This report contains the collective views of an international group of experts and does not necessarily represent the decisions or the stated policy of the United Nations Environment Programme, the International Labour Organization, or the World Health Organization.

Concise International Chemical Assessment Document 58

First draft prepared by Mr Peter Watts, Toxicology Advice & Consulting Ltd, Sutton, Surrey, United Kingdom; Mr G. Long, Health Canada, Ottawa, Canada; and Ms M.E. Meek, Health Canada, Ottawa, Canada

Published under the joint sponsorship of the United Nations Environment Programme, the International Labour Organization, and the World Health Organization, and produced within the framework of the Inter-Organization Programme for the Sound Management of Chemicals.

World Health Organization

Geneva, 2004

The International Programme on Chemical Safety (IPCS), established in 1980, is a joint venture of the United Nations Environment Programme (UNEP), the International Labour Organization (ILO), and the World Health Organization (WHO). The overall objectives of the IPCS are to establish the scientific basis for assessment of the risk to human health and the environment from exposure to chemicals, through international peer review processes, as a prerequisite for the promotion of chemical safety, and to provide technical assistance in strengthening national capacities for the sound management of chemicals.

The Inter-Organization Programme for the Sound Management of Chemicals (IOMC) was established in 1995 by UNEP, ILO, the Food and Agriculture Organization of the United Nations, WHO, the United Nations Industrial Development Organization, the United Nations Institute for Training and Research, and the Organisation for Economic Co-operation and Development (Participating Organizations), following recommendations made by the 1992 UN Conference on Environment and Development to strengthen cooperation and increase coordination in the field of chemical safety. The purpose of the IOMC is to promote coordination of the policies and activities pursued by the Participating Organizations, jointly or separately, to achieve the sound management of chemicals in relation to human health and the environment.

WHO Library Cataloguing-in-Publication Data

Chloroform.

(Concise international chemical assessment document ; 58)

1.Chloroform - adverse effects 2.Risk assessment 3.Environmental exposure

I.International Programme on Chemical Safety II.Series

ISBN 92 4 153058 8       (LC/NLM Classification: QV 81)

ISSN 1020-6167

©World Health Organization 2004

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Technically and linguistically edited by Marla Sheffer, Ottawa, Canada, and printed by Wissenchaftliche Verlagsgesellschaft mbH, Stuttgart, Germany

FOREWORD

Concise International Chemical Assessment Documents (CICADs) are the latest in a family of publications from the International Programme on Chemical Safety (IPCS) — a cooperative programme of the World Health Organization (WHO), the International Labour Organization (ILO), and the United Nations Environment Programme (UNEP). CICADs join the Environmental Health Criteria documents (EHCs) as authoritative documents on the risk assessment of chemicals.

International Chemical Safety Cards on the relevant chemical(s) are attached at the end of the CICAD, to provide the reader with concise information on the protection of human health and on emergency action. They are produced in a separate peer-reviewed procedure at IPCS. They may be complemented by information from IPCS Poison Information Monographs (PIM), similarly produced separately from the CICAD process.

CICADs are concise documents that provide summaries of the relevant scientific information concerning the potential effects of chemicals upon human health and/or the environment. They are usually based on selected national or regional evaluation documents or on existing EHCs. Before acceptance for publication as CICADs by IPCS, these documents undergo extensive peer review by internationally selected experts to ensure their completeness, accuracy in the way in which the original data are represented, and the validity of the conclusions drawn.

The primary objective of CICADs is characterization of hazard and dose-response from exposure to a chemical. CICADs are not a summary of all available data on a particular chemical; rather, they include only that information considered critical for characterization of the risk posed by the chemical. The critical studies are, however, presented in sufficient detail to support the conclusions drawn. For additional information, the reader should consult the identified source documents upon which the CICAD has been based.

Risks to human health and the environment will vary considerably depending upon the type and extent of exposure. Responsible authorities are strongly encouraged to characterize risk on the basis of locally measured or predicted exposure scenarios. To assist the reader, examples of exposure estimation and risk characterization are provided in CICADs, whenever possible. These examples cannot be considered as representing all possible exposure situations, but are provided as guidance only. The reader is referred to EHC 170.¹

While every effort is made to ensure that CICADs represent the current status of knowledge, new information is being developed constantly. Unless otherwise stated, CICADs are based on a search of the scientific literature to the date shown in the executive summary. In the event that a reader becomes aware of new information that would change the conclusions drawn in a CICAD, the reader is requested to contact IPCS to inform it of the new information.

Procedures

The flow chart on page 2 shows the procedures followed to produce a CICAD. These procedures are designed to take advantage of the expertise that exists around the world — expertise that is required to produce the high-quality evaluations of toxicological, exposure, and other data that are necessary for assessing risks to human health and/or the environment. The IPCS Risk Assessment Steering Group advises the Coordinator, IPCS, on the selection of chemicals for an IPCS risk assessment based on the following criteria:

• there is the probability of exposure; and/or
• there is significant toxicity/ecotoxicity.

Thus, it is typical of a priority chemical that

• it is of transboundary concern;
• it is of concern to a range of countries (developed, developing, and those with economies in transition) for possible risk management;
• there is significant international trade;
• it has high production volume;
• it has dispersive use.

The Steering Group will also advise IPCS on the appropriate form of the document (i.e., a standard CICAD or a de novo CICAD) and which institution bears the responsibility of the document production, as well as on the type and extent of the international peer review.

The first draft is usually based on an existing national, regional, or international review. When no appropriate source document is available, a CICAD may be produced de novo. Authors of the first draft are usually, but not necessarily, from the institution that developed the original review. A standard outline has been developed to encourage consistency in form. The first draft undergoes primary review by IPCS to ensure that it meets the specified criteria for CICADs.

The second stage involves international peer review by scientists known for their particular expertise and by scientists selected from an international roster compiled by IPCS through recommendations from IPCS national Contact Points and from IPCS Participating Institutions. Adequate time is allowed for the selected experts to undertake a thorough review. Authors are required to take reviewers’ comments into account and revise their draft, if necessary. The resulting second draft is submitted to a Final Review Board together with the reviewers’ comments. At any stage in the international review process, a consultative group may be necessary to address specific areas of the science. When a CICAD is prepared de novo, a consultative group is normally convened.

The CICAD Final Review Board has several important functions:

• to ensure that each CICAD has been subjected to an appropriate and thorough peer review;
• to verify that the peer reviewers’ comments have been addressed appropriately;
• to provide guidance to those responsible for the preparation of CICADs on how to resolve any remaining issues if, in the opinion of the Board, the author has not adequately addressed all comments of the reviewers; and
• to approve CICADs as international assessments.

Board members serve in their personal capacity, not as representatives of any organization, government, or industry. They are selected because of their expertise in human and environmental toxicology or because of their experience in the regulation of chemicals. Boards are chosen according to the range of expertise required for a meeting and the need for balanced geographic representation.

Board members, authors, reviewers, consultants, and advisers who participate in the preparation of a CICAD are required to declare any real or potential conflict of interest in relation to the subjects under discussion at any stage of the process. Representatives of nongovernmental organizations may be invited to observe the proceedings of the Final Review Board. Observers may participate in Board discussions only at the invitation of the Chairperson, and they may not participate in the final decision-making process.

This CICAD on chloroform was drafted by Toxicology Advice & Consulting Ltd based on documentation prepared by Environment Canada and Health Canada as part of the Priority Substances Program under the Canadian Environmental Protection Act (CEPA). The objective of assessments of priority substances under CEPA is to assess potential effects of indirect exposure in the general environment on human health as well as environmental effects. Data identified as of October 1999 were considered in the source document (Environment Canada & Health Canada, 2001). A comprehensive literature search of several on-line databases and other sources was conducted in February 2003 to identify any key references published subsequent to those incorporated in the source document. Information on the nature of the peer review and the availability of the source document is presented in Appendix 1. Information on the peer review of this CICAD is presented in Appendix 2. This CICAD was approved as an international assessment at a meeting of the Final Review Board, held in Varna, Bulgaria, on 8-11 September 2003. Participants at the Final Review Board meeting are listed in Appendix 3. The International Chemical Safety Card (ICSC 0027) for chloroform, produced by the International Programme on Chemical Safety (IPCS, 2000a), has also been reproduced in this document.

Chloroform (CAS No. 67-66-3) is a clear, colourless, volatile liquid with a pleasant etheric odour.

The total global flux of chloroform through the environment is approximately 660 000 tonnes per year, and about 90% of emissions are natural in origin. In the late 1990s, some 520 000 tonnes were manufactured annually, mainly in the USA, the European Union, and Japan. A major use is in the production of chlorodifluoromethane (HCFC-22), which is used (in decreasing quantities) as a refrigerant and (increasingly) as a fluoropolymer feedstock. Chloroform may be released into the environment from HCFC-22 plants. The other main chloroform releases to the environment occur as a result of using chlorine-based chemicals for bleaching and disinfection purposes at pulp and paper mills and water treatment plants.

Chloroform volatilizes readily from soil and surface water and undergoes degradation in air to produce phosgene, dichloromethane, formyl chloride, carbon monoxide, carbon dioxide, and hydrogen chloride. Its half-life in air ranges from 55 to 620 days. Biodegradation in water and soil is slow. Chloroform does not bioaccumulate to any significant extent in aquatic organisms. Chloroform is detected in outdoor air, usually at concentrations below 1 µg/m³. Indoor air concentrations can be approximately 10-fold higher, but may rise to about 1000 µg/m³ temporarily during hot-water showering in a poorly ventilated shower compartment. In drinking-water, mean chloroform concentrations of about 10-90 µg/litre have been reported in Canada. Mean total intake from food, drinking-water, and air was approximately 0.6-10 µg/kg body weight per day.

Chloroform is absorbed, metabolized, and eliminated rapidly by mammals following oral, inhalation, and dermal exposure. Oxidative metabolism, mainly CYP2E1 dependent, generates carbon dioxide as well as the toxic metabolites phosgene and hydrochloric acid. Metabolism of chloroform is much faster in mice than in humans.

Neat chloroform was irritating to human and rabbit eyes and to the skin of rabbits. Inhalation of chloroform causes anaesthesia in humans. Nasal lesions have also been observed in rats and mice exposed by inhalation or via the oral route. Laboratory animal studies identify the liver and kidneys as the key target organs of chloroform’s toxic potential, and limited data suggest that the liver and kidneys are the likely target organs in humans also. Informative epidemiological studies on chloroform were not identified. In laboratory animal bioassays, chloroform induced liver and kidney tumours. In rats, the only convincing evidence of carcinogenicity was an increase in kidney tumours in males given chloroform in a corn oil vehicle or in drinking-water. Kidney tumours were also seen in male mice exposed by inhalation or by ingestion in a toothpaste vehicle. In addition, male and female mice developed liver tumours when chloroform was delivered by gavage in a corn oil vehicle. Extensive investigation of chloroform’s genotoxicity potential generally failed to identify any activity, although some studies suggest that it may be weakly genotoxic in rats. A weight-of-evidence approach suggests that chloroform does not have significant genotoxic potential. There is convincing experimental evidence that the liver and kidney tumours seen in mice are a secondary consequence of sustained cytotoxicity (presumably due to metabolites such as phosgene and hydrogen chloride) and persistent associated reparative cell proliferation. Experimental support for a similar mechanism underlying the development of kidney tumours in male rats is more limited, but the data that are available are consistent with the proposed mechanism. Reproductive and developmental studies in a range of laboratory animal species suggest that chloroform is not a specific developmental toxin and is fetotoxic only at doses that cause maternal toxicity.

On repeated inhalation exposure, the lowest reported effect level in a laboratory animal study was 9.8 mg/m³, which caused cellular proliferation in nasal passage tissues of rats and mice. For repeated oral exposure, lowest reported effect levels were similar (10-17 mg/kg body weight per day) in various species for different end-points. A physiologically based pharmacokinetic (PBPK) model and the results from a 7.5-year dog study in which mild liver toxicity (fatty cysts suggestive of disruption of hepatic metabolism of fat) was seen were used to predict the rate of chloroform metabolism in the human liver (3.8 mg/litre per hour) that would produce a tissue dose rate of toxic metabolites associated with a 5% increase in risk. This tissue dose rate would result from lifetime drinking of water containing chloroform at 37 mg/litre or lifetime exposure to 9.8 mg chloroform/m³ air. Respective lower 95% confidence limits were 12 mg/litre and 3.4 mg/m³. A tolerable daily oral intake of 0.015 mg/kg body weight per day and a tolerable concentration of 0.14 mg/m³ air are derived from these figures.

In addition, the PBPK model and the results from a study in which chloroform induced kidney tumours in male rats were used to derive analogous human rates of metabolism leading to a 5% increase in the incidence of tumours and tumour precursor lesions. These were estimated to be 3.9 and 1.7 mg/litre per hour, respectively. For the former, the 95% lower confidence limits for continuous exposure via drinking-water and via air were 2363 mg/litre and 74 mg/m³, respectively. For the latter, the metabolic rate was equivalent to continuous exposure at 1477 mg/litre water and 33.3 mg/m³ air (95% lower confidence limits were not given).

In a sample risk characterization, the margins between estimated exposure of the general population in Canada and tumorigenic and benchmark doses for cancer and non-cancer effects, respectively, for chloroform were greater than 2 orders of magnitude.

The lowest concentration reported to cause cellular proliferation in the nasal cavities of rats and mice (9.8 mg/m³) is 4298 and 1225 times higher, respectively, than the midpoint (2.28 µg/m³) and 95th percentile (8.0 µg/m³) estimates for chloroform in indoor air in Canada.

No toxicity data were identified for birds or wild mammals, but laboratory animal data indicate that atmospheric emissions of chloroform do not pose any significant risks to terrestrial wildlife. No directly relevant data were available for estimating potentially harmful concentrations in soil. For aquatic organisms, concentrations in surface waters are rarely above estimated toxicity thresholds, even for sensitive species. There is some uncertainty regarding exposure levels — and hence possible risks to aquatic organisms — near industrial leachate sources such as pulp and paper mills, water treatment plants, and landfill sites.

Chloroform (CAS No. 67-66-3) is also known as trichloromethane, methane trichloride, trichloroform, methyl trichloride, and formyl trichloride. Its molecular formula is CHCl₃, and its relative molecular mass is 119.4. Chloroform’s chemical structure is shown in Figure 1.

Fig. 1: Chemical structure of chloroform

At room temperature, chloroform is a clear, colourless, volatile liquid with a pleasant etheric odour. The ranges of values reported for selected physical/chemical properties are presented in Table 1. Additional properties are given in the International Chemical Safety Card (ICSC 0027) reproduced in this document.

Table 1: Physical and chemical properties of chloroform.

^a Data listed in source document (Environment Canada & Health Canada, 2001).

The conversion factors² for chloroform in air at 20 °C and 101.3 kPa are as follows:

1 ppm = 4.96 mg/m³

1 mg/m³ = 0.202 ppm

In this CICAD, we have followed the convention of the source document, which is to use conversion factors at 25 °C instead of 20 °C:

1 ppm = 4.9 mg/m³

1 mg/m³ = 0.204 ppm

The general method of quantifying chloroform in water samples involves preservation of samples with sodium thiosulfate, without prior pH adjustment, and analysis by gas chromatography (GC) with electron capture detection (ECD), halogen-specific detection, or mass-selective detection. There are two recommended International Organization for Standardization methods (ISO, 1997). The first involves liquid-liquid extraction GC with ECD or other suitable detector. Pentane, hexane, petroleum ether, heptane, or xylene (for wastewater) are used as extraction solvents, and the quantification limits are 0.05-0.3 µg/litre. The second method involves static headspace GC with ECD or other suitable detector and has a quantification limit of 0.3 µg/litre. The recommended US Environmental Protection Agency (EPA) methods involve purge and trap GC with electrolytic conductivity or microcoulometric detectors (EPA Method 502) or purge and trap GC-mass spectrometry (MS) (EPA Method 524). Quantification limits are 0.02-0.2 µg/litre.

A number of analytical methods may be used to determine chloroform concentrations in air. The most common procedures use GC techniques with ECD, flame ionization detection, photoionization detection, or MS. Chloroform can be measured directly in a procedure in which air is aspirated or injected directly into the measuring instrument without pretreatment. Although these methods are simple, they can be used only when chloroform is present in the air at relatively high levels (e.g., urban source areas). In a second major method (adsorption-liquid desorption), air samples are passed through an activated adsorbing agent (e.g., charcoal or Porapak-N). The adsorbed chloroform is then desorbed with an appropriate solvent (e.g., carbon disulfide or methanol) and subsequently passed through the GC for measurement. In the adsorption-thermal desorption technique, air samples are also passed through an activated absorbing agent (e.g., Tenax-GC, Porapak-Q, Porapak-N, or carbon molecular sieve). The adsorbed chloroform is then thermally desorbed and driven into the GC column for determination. The fourth major technique (cold trap-heating) involves injection of air samples into a cold trap (liquid nitrogen or liquid oxygen is used for cooling). The trap is then heated while transferring its chloroform content into the column of a GC for measurement. Details on currently used methods may be obtained from the US Occupational Safety and Health Administration (OSHA), United Kingdom Health and Safety Executive, American Society for Testing and Materials (ASTM), US National Institute for Occupational Safety and Health (NIOSH), and US EPA.³

The sensitivity of analytical methods has improved over time; lowest detection limits reported in the source document are 0.1 µg/m³ in air (T. Dann, personal communication, 1998), 0.001 µg/litre in water (Comba et al., 1993), 0.05 µg/kg in dry food (Page & Lacroix, 1993), and 0.02 µg/kg in beverages (McNeal et al., 1995).

Based on estimated half-time and measured concentrations in different parts of the world, the total release of chloroform to the air was estimated to be 470 000 tonnes per year (Khalil & Rasmussen, 1999). A review paper published in 2003 reported that the chloroform flux through the environment is apparently constant at some 660 000 tonnes per year and that about 90% of emissions are natural in origin. This global flux consisted of 360 000 ± 90 000, 220 000 ± 100 000, <20 000, and 66 000 ± 23 000 tonnes per year from offshore seawater, soil processes, other natural sources (including volcanic activity and geological), and anthropogenic activities, respectively (McCulloch, 2003).

The natural production of chloroform by marine macroalgae has been reported (Nightingale et al., 1995; Scarratt & Moore, 1999). The release of chloroform (12 µg/m² per day) from an organic-rich spruce forest soil, under aerobic conditions in the laboratory, suggested biogenic formation (Haselmann et al., 2000). Chloroform was found at 20-30 µg/m³ in soil air down to a depth of 160 cm, compared with a reported value of about 0.1 µg/m³ in the atmosphere. Soil spiking studies with radiolabelled chloride (Na³⁷Cl) demonstrated natural formation in soil. Fungi were believed to play an important role in this natural production (Hoekstra et al., 1998).

One group has suggested that natural and anthropogenic sources make approximately equal contributions to atmospheric chloroform. This group measured chloroform flux at seven peatland locations and two evergreen forested bog sites in Ireland in 1998 and estimated annual global fluxes of 4700 (range 100-151 900) tonnes per year from peatland ecosystems and 24 100 (range not given) tonnes per year from total wetlands (Dimmer et al., 2001). As mentioned above, McCulloch (2003) reported an approximate global chloroform flux of 660 000 tonnes per year and estimated that about 90% of these emissions were natural in origin.

Chloroform can be released to the environment from direct processes (production, storage, transit, or use) or as a result of its formation from other substances, in processes such as paper bleaching with chlorine and water chlorination. Pulp and paper mills, municipal wastewater treatment plants, chemical manufacturing plants, and waste incinerators represent anthropogenic sources of chloroform (IPCS, 1994a). Various organic compounds present in natural waters, particularly humic and fulvic acids derived from soils and the decomposition of plant material, may contribute to the formation of chloroform (via the "haloform reaction") in areas where the drinking-water has been chlorinated (Environment Canada & Health Canada, 2001). As mentioned above, McCulloch (2003) reported that anthropogenic sources contribute about 66 000 ± 23 000 tonnes per year.

An industrial survey carried out in Canada revealed reported releases, by 23 pulp and paper mills, of 288 tonnes of chloroform into the atmosphere, 15.6 tonnes into water bodies, 0.019 tonnes into wastewater treatment plants, and 0.127 tonnes into landfills in 1996 (Environment Canada, 1997a). Chloroform generation and concentrations in effluents of these mills are reduced significantly when chlorine dioxide is substituted for elemental chlorine in the bleaching process (Solomon et al., 1994; M. Henteleff, personal communication to Environment Canada, 1999).

In 1996, total on-site environmental releases of chloroform reported to the Canadian National Pollutant Release Inventory were 208 tonnes. Almost all was released by the pulp, paper, and allied products industry; more than 96% was released to the atmosphere, with the remainder being released to water (NPRI, 1999).

Although not quantified, Canadian municipal wastewater treatment plant disinfection systems that use chlorine can be significant sources of chloroform. Chloroform is produced by the reaction between chlorine and organic precursor molecules such as fulvic and humic acids (Environment Canada, 1999a; Environment Canada & Health Canada, 2001).

Chloroform can also be released from industrial plants. A Canadian survey revealed that three facilities belonging to members of the Canadian Chemical Producers’ Association released a total of 145 kg chloroform in 1996, of which 88% was released to air (Environment Canada, 1997a). The Canadian Chemical Producers’ Association estimated that its member companies released 540 kg to the environment in 1992 (CCPA, 1992). In 1993, chloroform releases as a result of its use in HCFC-22 production were estimated to range from 31 to 1040 kg (Environment Canada & Health Canada, 2001).

Chloroform is manufactured mainly in the USA, the European Union, and Japan, the total global capacity in the late 1990s being about 520 000 tonnes per year (McCulloch, 2003). In 1995, chloroform was produced in 19 countries. The volume of production of chloroform in the USA was 229 000 tonnes in 1991 and 216 000 tonnes in 1993 (IARC, 1999). Chloroform is no longer produced in Canada (Environment Canada & Health Canada, 2001). The total production in the European Union has been estimated at 316 000 tonnes (ECSA, 1997).

Chloroform’s main use is in HCFC-22 production, and this accounts for 90-95% of its use in the European Union (Zok et al., 1998). Although use of HCFC-22 in refrigerant applications is decreasing, increasing use of HCFC-22 as the feedstock for fluoropolymers such as polytetrafluoroethylene means that demand for chloroform has remained relatively constant. Earlier use of chloroform as an anaesthetic has been largely discontinued in Canada, but it still has limited use in some dental procedures and in certain pharmaceuticals. The Montreal Protocol, as amended, means that HCFC-22 will be phased out between 2010 and 2020, effectively eliminating much of the present market for chloroform (Environment Canada & Health Canada, 2001). Worldwide, chloroform is also used in pesticide formulations, as a solvent for fats, oils, rubber, alkaloids, waxes, gutta-percha, and resins, as a cleansing agent, in fire extinguishers, and in the rubber industry (ESCA, 1997; Budavari, 2001).

Chloroform emitted to air reacts primarily with photochemically generated hydroxyl radicals in the troposphere (Kindler et al., 1995). Reaction products include phosgene, dichloromethane, formyl chloride, carbon monoxide, carbon dioxide, and hydrogen chloride (Gürtler & Kleinermanns, 1994). Experimentally derived rate constants for this reaction at 25 °C range from 1.0 × 10^-13 to 2.95 × 10^-13 cm³/molecule per second. Its rate of decomposition depends on a number of factors, including temperature, hydroxyl radical concentration, and the number of hours of sunshine. Estimated half-lives vary between about 55 and 620 days (Derwent & Eggleton, 1978; Singh et al., 1981; Klöpffer et al., 1988; Khalil & Rasmussen, 1999). Wet deposition is considered minor, as most will return to the air by volatilization (Diamond et al., 1994). Mass destruction rates have been estimated to be 250 000-570 000 and 120 000-260 000 tonnes per year in the northern and southern hemispheres, respectively (McCulloch, 2003).

In surface water, the principal removal process is volatilization. Modelling studies have generated estimated half-lives of 1.5 days and 9-10 days in a river and a lake, respectively (US EPA, 1984). Other models have indicated shorter half-lives in shallow, well-mixed systems with high wind velocities (Kaczmar, 1979; Lyman et al., 1982). Most studies have indicated little biodegradation after up to 25 weeks in aquatic systems under aerobic conditions (Bouwer et al., 1981; Wilson et al., 1981, 1983; Bouwer & McCarty, 1984). In groundwater, restricted volatilization and slow biodegradation (under anaerobic conditions) or no biodegradation (under most aerobic conditions) means that chloroform may be quite persistent (Environment Canada & Health Canada, 2001). The half-life by hydrolysis has been reported to be greater than 1000 years (McCulloch, 2003).

Limited studies suggest that chemical degradation in sediments is not rapid, except under anaerobic methanogenic conditions. The major degradation products under anaerobic conditions are carbon dioxide, methane, and hydrogen chloride, with smaller amounts of dichloromethane. Under anaerobic conditions, chloroform had half-lives of 12 days at 10 °C and 2.6 days at 20 °C (Van Beelen & Van Keulen, 1990). In another study carried out under anaerobic conditions, chloroform was degraded in muddy sediments with a half-life of 2-37 days, whereas no degradation could be demonstrated in sandy sediments (Van Beelen & Van Vlaardingen, 1993).

The principal fate of chloroform at the soil surface is temperature-dependent volatilization, due to its volatile nature and low soil adsorption. A microcosm study involving daily application of wastewater containing chloroform to soil found that 75% of applied chloroform volatilized to the air, while the remainder leached through the soil (Piwoni et al., 1986). Chloroform adsorption is correlated with soil clay content (Dural & Peng, 1995). Limited studies suggest that chemical degradation in soil is not rapid, except under anaerobic methanogenic conditions. The major degradation products under anaerobic conditions are carbon dioxide, methane, and hydrogen chloride, with smaller amounts of dichloromethane (Van Beelen & Van Keulen, 1990; Van Beelen & Van Vlaardingen, 1993).

The octanol/water partition coefficient (log K_ow = 1.97) indicates that chloroform is unlikely to bioaccumulate to any significant extent in aquatic biota (Anderson & Lusty, 1980; Zok et al., 1998). Reported bioconcentration factors include 690 in green algae (Mailhot, 1987) and 1.4-10 in various fish (bluegill Lepomis macrochirus, rainbow trout Oncorhynchus mykiss, largemouth bass Micropterus salmoides, and channel catfish Ictalurus punctatus) (Veith et al., 1978; Anderson & Lusty, 1980; Barrows et al., 1980). Depuration is rapid, with a half-life of less than 1 day in all of the above fish species (Anderson & Lusty, 1980; Barrows et al., 1980).

Chloroform in soil or surface water volatilizes readily; at equilibrium, greater than 99% is expected to partition to the atmosphere (Zok et al., 1998; McCulloch, 2003). Due to chloroform’s water solubility, some wet deposition of atmospheric chloroform may occur, but subsequent revolatilization is likely to be extensive (Diamond et al., 1994). Chloroform is not expected to partition significantly to soils or sediments, because its affinity for organic carbon and lipids is low. Modelling has predicted that the percentage of chloroform in water transferred to bottom sediments would range from <0.06% in lakes to 8% in ponds (Anderson et al., 1985). Compartmental partitioning has been reported to be 99.1%, 0.9%, 0.01%, and 0.01% in air, water, soil, and sediment, respectively (Zok et al., 1998).

6.1.1 Ambient air

Over land, there is substantial variability in chloroform concentration. Mean levels in urban and/or industrial areas ranged up to 3.5 µg/m³, with most concentrations in the range 0.5-1.5 µg/m³ (McCulloch, 2003). Median values in air from Madeira, the Portugese coast, and the Black Forest in a 1991 report were 0.2-0.6 µg/m³ (range 0.07-8.7 µg/m³), and measurements reported in a 1975 paper quantified chloroform at 0.12-0.6 µg/m³ in rural air in the United Kingdom (McCulloch, 2003).

Chloroform was detected (i.e., above the detection limit of 0.1 µg/m³) in approximately 69% of 8807 24-h samples collected from 47 sites in seven Canadian provinces between 1989 and 1996. During this period, annual median and mean concentrations ranged from <0.1 to 0.18 µg/m³ and from 0.12 to 0.23 µg/m³, respectively. Concentrations were lowest at rural sites, higher at urban sites, and highest immediately adjacent to major roadways. Comparison of 3344 samples taken during 1989-1992 with 5463 samples taken between 1993 and 1996 indicated that chloroform concentrations were slightly lower in the more recent period. The highest 24-h average concentration detected before 1996 was 6.0 µg/m³, compared with 0.75 µg/m³ during 1996 (T. Dann, personal communication, 1998).

Atmospheric chloroform levels of 0.1-10 µg/m³ and 1.4-110 µg/m³ have been reported for urban and source-dominated areas in the USA (ATSDR, 1996).

Based on a 9-year series of measurements at the surface of the polar, middle, and tropical latitudes of both hemispheres, an average surface concentration of 0.09 µg/m³ was reported, although annual averages at some continental locations rose to 0.2 µg/m³. No significant trends were seen during the period of measurement (Khalil & Rasmussen, 1999).

Chloroform levels in the air over the southern Atlantic Ocean were reported to be 0.05-0.1 µg/m³. Equivalent figures for the northern hemisphere were 0.1-0.25 µg/m³, with lower levels of 0.04-0.07 µg/m³ above the trade wind system. These figures were based on lower troposphere samples taken aboard a ship cruising from Capetown to Bremerhaven during 1985, together with samples taken in the Azores in 1982, in Madeira in 1984, and in Bermuda in 1985 (Class & Ballschmiter, 1986).

6.1.2 Indoor air

Chloroform was detected (detection limit 3.5 µg/m³) in 11% of 24-h samples taken in 754 residences in nine Canadian provinces during 1991. The maximum concentration found was 68.6 µg/m³ (Concord Environmental Corporation, 1992; Health Canada, 1999). Chloroform was detected (detection limit 2.3 µg/m³) in 8 of 44 households in Ontario (Greater Toronto Area), Nova Scotia, and Alberta during 1996 and in 34 of 50 households (detection limit 0.22 µg/m³) in the same areas during 1997. The maximum concentration detected was 14.1 µg/m³, and the estimated mean concentration for the total of 94 samples (assuming chloroform was present at half of the appropriate detection limit for each "non-detect" sample) was 1.5 µg/m³. Single personal breathing zone samples taken from all 94 households had concentrations ranging from undetected (<0.22 µg/m³) to 94.5 µg/m³, with an overall estimated mean of 2.6 µg/m³ (Otson & Meek, 1996; Conor Pacific Environmental, 1998; Health Canada, 1999). Indoor air sampling found chloroform in 89 of 146 households in Windsor, Ontario, during 1991-1992 (detection limit unknown). In the indoor air of non-smoking households, the highest mean concentration was 5.6 µg/m³; the concentration was higher (16 µg/m³) where environmental tobacco smoke was present (OMEE, 1994). However, no differences in concentrations were seen in the air of 61 non-smoking households and 32 smoking households in New Jersey, USA, in 1962 (means 0.60 and 0.85 µg/m³, respectively; medians 0.28 and 0.23 µg/m³, respectively) (Heavner et al., 1996), and tobacco was not identified by IPCS (1994a) as an important source of environmental chloroform exposure. Mean concentrations of 0.17-43.9 µg/m³ (maximum 210 µg/m³) have been reported for indoor air in the USA (Samfield, 1992). Mean concentrations in 248 homes in Los Angeles, California, USA, in 1987 ranged from 0.9 to 1.5 µg/m³ (maximum 13 µg/m³) (Wallace, 1997).

Indoor air concentrations of chloroform may rise for short periods of time due to volatilization from hot water. In particular, chloroform concentrations in the shower compartment while taking a shower may exceed 1000 µg/m³, due to volatilization of more than 50% of dissolved chloroform (Tancrède et al., 1992; Giardino & Andelman, 1996; Health Canada, 1999).

6.1.3 Surface water

Chloroform concentrations ranging from 0.002 to 0.015 µg/litre have been reported for the open ocean (Class & Ballschmiter, 1986; IPCS, 1994a; Zok et al., 1998). A review reported chloroform concentrations in North Sea coastal water and estuarine waters of several European countries (including France, Germany, Sweden, and the United Kingdom) in the 1980s to 1990s to range from 0.004 to 11.5 µg/litre. Typical background levels in rivers in non-industrialized areas were generally below 0.5 µg/litre, while levels of up to 10 µg/litre were detected in rivers in industrialized areas or in the vicinity of emission points from municipal wastewater treatment plants (Zok et al., 1998). Chloroform concentrations ranging from <0.01 to 70 µg/litre have been reported for European estuaries, with locally high levels related to point source discharges (McCulloch, 2003).

Analysis of 59 samples of surface water and well water in Alberta, Canada, during 1990-1995 revealed only two samples (containing 2 and 7 µg/litre) above the 1 µg/litre detection limit (Alberta Environment, 1996). Chloroform was detected (detection limit 1 µg/litre) in only a few of 321 samples of surface water from Alberta in 1990-1996, with the highest concentration reported as 2 µg/litre (Environment Canada, 1996). In water from Lake Superior, chloroform levels varied from <0.001 to 4.2 µg/litre (median 0.064 µg/litre) in 192 samples taken during 1991 (Comba et al., 1993), and a maximum level of 0.19 µg/litre was found in 293 samples taken from the Niagara River during 1990-1993 (Environment Canada, 1996). Concentrations in 107 samples of Quebec surface waters collected from 1990 to 1993 ranged from non-detectable (<0.2 µg/litre) to 44 µg/litre (MENVIQ, 1996). Across four Canadian provinces, the median value (of 984 measurements) was <0.2 µg/litre, and the 95th and 99th percentiles were <1 and 2.94 µg/litre, respectively (Environment Canada & Health Canada, 2001).

High concentrations have occasionally been reported in Canadian surface waters, particularly near pulp and paper mills using chlorine bleaching (e.g., chloroform concentrations below a mill in 1986 ranged from 80 to 200 µg/litre) (OMOE, 1990). Similarly, concentrations of up to 394 µg/litre were reported in rivers sampled in the 1970s in highly industrial US cities (IPCS, 1994a).

Rainwater concentrations of 11-17 ng/litre and up to 97 ng/litre were reported in two studies carried out in the Black Forest area in the 1990s (McCulloch, 2003).

6.1.4 Drinking-water

Chloroform is the principal by-product of water disinfection processes such as chlorine-chloramine, chlorine-chlorine, and ozone-chlorine treatment. Levels vary widely depending upon concentrations of organic materials in the raw water and are also influenced by treatment method, temperature, and pH; chloroform concentrations increase in summer months, as the water moves along the distribution system, and in domestic hot water tanks (Environment Canada & Health Canada, 2001). The majority of data collected in Canada originate from measurements collected within water treatment plants and distribution systems; little information on levels at the consumer tap is available. Concentrations measured (detection limits 0.1-1.0 µg/litre) in drinking-water in several areas in Canada during the 1990s are presented in Table 2. The data were used to derive a 95th-percentile value of 166 µg/litre. Using only data from the two areas of highest concentrations (to establish a "reasonable worst case"), the 95th-percentile value was 220 µg/litre (Health Canada, 1999).

Table 2: Concentrations of chloroform in drinking-water in Canada during the 1990s.

More limited data are available from smaller national studies. In 1993, median, mean, and maximum concentrations of 13.4, 27.6, and 336 µg/litre, respectively, were recorded in 214 samples collected from 53 water treatment facilities in nine Canadian provinces; chloroform was detected (>0.2 µg/litre) in all samples. Arithmetic means among the provinces varied from 6.5 to 62.1 µg/litre and were about twice as high in summer as in winter (Williams et al., 1995; Health Canada, 1999).

Chloroform was measured at 24 µg/litre in warm water entering a shower when the cold water contained about 6 µg/litre in the winter or 12 µg/litre in the summer (Benoit et al., 1997).

In a Canadian study carried out in 1992, chloroform (detection limits ranged from 0.5 to 3.0 µg/litre) was not found in any of 61 bottles of mineral water and was detected (at 3.7 µg/litre) in only 1 of 86 spring water samples. Chloroform was detected in 10 of 35 further samples of bottled water, including carbonated, demineralized, deionized, treated, and distilled waters (Page et al., 1993).

6.1.5 Sediment and soil

Although only limited data were identified, chloroform does not appear to be sorbed in sediments or soils to any great extent, and so it is unlikely that it will accumulate in these media to any significant extent (Environment Canada, 1999a; Environment Canada & Health Canada, 2001).

6.1.6 Food

The sources of chloroform in food are not clearly understood, although migration of chloroform from packaging solvents, glues, and inks has been documented, and transfer from surfaces cleaned with chlorinated water to lipid-containing foods contacting these surfaces is a possibility. The use of chlorinated water by bottling plants (e.g., soft drink manufacturers) may explain the presence of chloroform in some beverages. Chloroform introduced to foods as a consequence of the use of chlorinated drinking-water during food preparation probably escapes by volatilization during cooking, reducing the concentrations in the table-ready foods (Environment Canada & Health Canada, 2001).

Chloroform was detected (at up to 14.8 µg/kg) in 11 of 13 beverages purchased in Ottawa, Ontario, Canada, but not in dry foods (detection limit 0.05 µg/kg). Subsequent sampling of 47 foods and beverages found chloroform in 41 samples at concentrations ranging from 0.23 to 129 µg/kg. The highest three concentrations (50-129 µg/kg) were found in butter (Page & Lacroix, 1993).

Analysis of composite food groups prepared from groceries bought from four retailers in Windsor, Ontario, Canada, found chloroform in 5 of 33 composites (cheese/butter, canned meats, vine vegetables, soft drinks, and dehydrated soups). The maximum concentration found was 67 µg/litre (Enviro-Test Laboratories, 1992). A similar study of 35 composite groups found chloroform (detection limits were 1 µg/litre in beverages and 5 µg/kg in foods) only in soft drinks and alcoholic drinks (Enviro-Test Laboratories, 1993).

Limited data are also available from the USA. Chloroform was found in 94 of 231 table-ready foods obtained from the US Food and Drug Administration’s (FDA) market basket collection, with the highest concentration (312 µg/kg) found in cheddar cheese (Daft, 1988). In an analysis of grain-based products, chloroform concentrations ranged from 0.5 µg/kg in lasagna noodles to 3400 µg/kg in wheat (Heikes & Hopper, 1986). Chloroform was found in 10 of 18 table-ready food samples; the highest concentration was 670 µg/kg in butter (Heikes, 1987). Chloroform was present at 30-255 µg/kg in 36 butter samples collected in Washington, DC (Miller & Uhler, 1988). Analysis of 234 foods revealed chloroform (detection limit 5 µg/kg) in 44 samples, including margarine (7.3 µg/kg), butter (38.9 µg/kg), and cream cheese (110 µg/kg) (Heikes et al., 1995).

Chloroform was detected at 0.1-65 µg/litre in 40 of 42 breast milk samples from nursing mothers in five US hospitals (Erickson et al., 1980).

Deterministic estimates of average and upper-bounding estimates for daily intake have been developed in Canada based on concentrations determined in Canadian air (national surveys), food in Canada and the USA, and drinking-water (provincial and territorial data) (Environment Canada & Health Canada, 2001). These are presented in Tables 3 and 4.

Table 3: Deterministic estimates of average daily intakes for the general population.^a

^a Further details on the basis for estimated figures are given in Environment Canada & Health Canada (2001).

^b Inhalation and dermal intake from daily showering.

Table 4: Deterministic upper-bounding estimates of daily intake for the general population.^a

^a Further details on the basis for estimated figures are given in Environment Canada & Health Canada (2001).

^b Inhalation and dermal intake from daily showering.

Deterministic estimates were generated using the above monitoring data and reference values for body weight, inhalation volume, and consumption of food and water. Average intake from food, drinking-water, and air varied from 0.6 to 10 µg/kg body weight per day. Upper-bounding estimates were calculated using maximum reported concentrations in water, food, and air and ranged from 40 to 95 µg/kg body weight per day (or up to 148 µg/kg body weight per day for infants fed exclusively on powdered infant formula prepared with tap water containing the maximum reported chloroform concentration). Daily showering increased estimated exposure by about 50-100% for some subgroups. Further details are given in the source document (Environment Canada & Health Canada, 2001).

In addition, probabilistic estimates of daily chloroform intake from air and drinking-water in Canada were developed for two scenarios (average population exposure and reasonable worst case), but data were considered insufficient to develop probabilistic exposure estimates from food consumption or showering. Simulations of 10 000 trials were run 5 times each using Monte Carlo random and Latin Hypercube methods. The two sampling methods gave similar estimates, and relative standard deviations (for n = 5 simulations of 10 000 trials each) of the upper-percentile estimates of intake did not exceed 5%, indicating a high degree of reproducibility. The average population scenario was based on the distribution of chloroform in 8807 outdoor air samples collected during the 1990s, the estimated geometric mean and standard deviation of an assumed lognormal distribution of chloroform in the indoor air of 754 Canadian homes, and analysis of chloroform in 6607 drinking-water samples in Canadian provinces and territories. The 95th percentiles of the distribution of intakes from inhalation and ingestion of drinking-water for five age groups of the general population (i.e., 0.5 years to 60+ years of age) ranged from 4.9 to 12.9 µg/kg body weight per day (Health Canada, 1999). The limitations of the data on the daily intake rate of total tap water by infants (EHD, 1998) prevented the development of probabilistic estimates for this subgroup.

The reasonable worst-case scenario was based on 800 outdoor air samples collected during the 1990s from four sites adjacent to major Canadian roadways, the estimated geometric mean and standard deviation of an assumed lognormal distribution of chloroform in the indoor air of 754 Canadian homes, and the distribution of chloroform in 2597 drinking-water samples from the two provinces with the highest reported concentrations. The 95th percentiles of the distribution of intakes from inhalation and ingestion of drinking-water for the same five age groups of the general population ranged from 7.0 to 19.1 µg/kg body weight per day (Health Canada, 1999). The limitations of the data on the daily intake rate of total tap water by infants (EHD, 1998) prevented the development of probabilistic estimates for this subgroup.

Chloroform was found (detection limit 0.1 µg/litre) in 54% of 979 samples of human blood collected in the USA, but concentrations were not quantified (Ashley et al., 1994). Concentrations in the urine of healthy male students in New Jersey, USA, ranged from 36.5 to 48.7 µg/litre (Youssefi et al., 1978).

The HSDB (2003) chloroform record includes a brief mention of mean time-weighted average (TWA) exposures of 13, 2, and 1 mg/m³ for production operators, drummers/bottle fillers, and maintenance/utility personnel at the Shell Chemical Company, Rocky Mountain Arsenal (a pesticide plant), levels of 10-1000 mg/m³ in a Polish pharmaceutical plant, an 8-h TWA of 77.4 mg/m³ (range 13-227 mg/m³) in a police forensic laboratory, and (during 1968-1972) levels of 34-830 mg/m³ (mean 230 mg/m³, 79 samples) in a film manufacturing plant using a solvent containing 22% chloroform (Santodonato et al., 1985).

Chloroform is well absorbed, metabolized, and eliminated rapidly by mammals after oral, inhalation, or dermal exposure (IPCS, 2000b). In humans given a single oral dose of 0.5 g chloroform, about 50-52% of the dose was absorbed, and virtually all of the absorbed dose was metabolized to carbon dioxide. Blood levels peaked after 1.5 h and declined in line with a two-compartment model with half-lives of 13 and 90 min, respectively (Fry et al., 1972). Following a single inhalation exposure to approximately 5 mg [³⁸Cl]chloroform, volunteers absorbed about 80% (Morgan et al., 1970). The relative contributions of dermal and pulmonary uptake have been studied in individuals taking showers, using post-exposure exhaled air concentrations to estimate uptake. These were 6-21 µg/m³ for normal showering and 2.4-10 µg/m³ if exposure during showering was restricted to the inhalation route ("inhalation-only" showers). This difference was statistically significant and indicated that the contributions of dermal and inhalation exposures were approximately equivalent (Jo et al., 1990).

Species differences can be seen. When rats, mice, and monkeys were given radiolabelled chloroform at 60 mg/kg body weight by the oral route, about 90% was absorbed and exhaled in all three species in the 48 h following dosing. However, while mice excreted about 85% of the dose as exhaled carbon dioxide and 5% as unchanged chloroform, monkeys exhaled only 18% as carbon dioxide and 79% as chloroform. The rat was intermediate, with 67% exhaled as carbon dioxide and 20% as chloroform. Excretion in the urine/faeces combined accounted for only about 2-3% of the dose in mice and monkeys and about 8% in rats (D.M. Brown et al., 1974). Metabolism of chloroform is much faster in mice than in humans. For example, the mean peak rate of metabolism at an inhalation exposure of 49 mg/m³ has been predicted to be approximately 78 times lower in humans than in mice (Delic et al., 2000).

Corley et al. (1990) measured radioactivity in the exhaled air, urine, faeces, carcass and skin, and cage wash in the 48 h following a 6-h inhalation exposure of rats and mice at various chloroform concentrations (49, 440, and 1790 mg/m³ for mice; 460, 1740, and 5100 mg/m³ for rats). At the low concentration, metabolism was extensive in both species. In mice, exhaled carbon dioxide, exhaled chloroform, urine, and faeces accounted for 7.22, 0.03, 0.95, and 0.05 mg equivalents/kg body weight, respectively; in rats, these figures were 31.84, 0.76, 3.34, and 0.04, respectively. However, partial saturation of metabolism was indicated at about 1800 mg/m³; in mice, the equivalent figures were 217.85, 23.03, 21.24, and 3.84 mg equivalents/kg body weight, respectively, while in rats, the equivalent figures were 54.85, 16.15, 6.53, and 0.81 mg equivalents/kg body weight, respectively (Corley et al., 1990).

Following a 10-min inhalation exposure of mice to [¹⁴C]chloroform (dose 280 mg/kg body weight), whole-body autoradiography carried out immediately after exposure or 2 h later showed high concentrations in the fat, blood, lungs, liver, kidneys, spinal cord and nerves, meninges, and cerebellar cortex. Non-volatile radioactivity was bound in the bronchi, nasal mucosa, liver, kidneys, salivary glands, and duodenal contents. High levels of volatile or extractable radioactivity were found in testes, preputial gland, and epididymis (Bergman, 1984). Transplacental transfer has been demonstrated in rats, mice, and guinea-pigs (Nicloux, 1906; Withey & Karpinski, 1985; Danielsson et al., 1986).

Both oxidative and reductive pathways of chloroform metabolism have been identified, although data in vivo are limited. Carbon dioxide is the major metabolite of chloroform generated by the oxidative pathway in vivo. The oxidative pathway also generates reactive metabolites, including phosgene (Pohl et al., 1977; Pohl & Krishna, 1978) (determined in vitro, with phenobarbital induction), while the reductive pathway generates the dichloromethylcarbene free radical (Wolf et al., 1977; Tomasi et al., 1985; Testai & Vittozzi, 1986) (determined in vitro and in vivo, both with and without phenobarbital induction). Oxidative and reductive metabolism both proceed through a cytochrome P450 (CYP)-dependent enzymatic activation step. The balance between oxidative and reductive pathways depends on species, tissue, dose, and oxygen tension. In intact mammals, oxidative tension probably precludes any significant metabolism by the reductive pathway (Testai & Vittozzi, 1986; Ammann et al., 1998). Phosgene is produced by oxidative dechlorination of chloroform to trichloromethanol, which spontaneously dehydrochlorinates (Mansuy et al., 1977; Pohl et al., 1977). Dehydrochlorination of trichloromethanol produces one molecule of hydrochloric acid, and hydrolysis of phosgene produces another two molecules, so that three molecules of hydrochloric acid are produced in the conversion of chloroform to carbon dioxide.

The electrophilic metabolite phosgene binds covalently to nucleophilic components of tissue proteins (Pohl et al., 1980). It also interacts with other cellular nucleophiles (Uehleke & Werner, 1975) and binds to some extent to the polar heads of phospholipids (Vittozzi et al., 1991). Alternatively, phosgene reacts with water to release carbon dioxide and hydrochloric acid (Fry et al., 1972; B.R. Brown et al., 1974; D.M. Brown et al., 1974). The interaction of phosgene with glutathione results in the formation of S-chlorocarbonyl glutathione, which can either interact with an additional glutathione to form diglutathionyl dithiocarbonate (Pohl et al., 1981) or form glutathione disulfide and carbon monoxide (Ahmed et al., 1977; Anders et al., 1978). Incubation of mouse renal microsomes with glutathione increases production of these metabolites from chloroform and decreases irreversible binding to proteins and further metabolism to carbon dioxide (Smith & Hook, 1984). Reduced glutathione is capable of scavenging essentially all chloroform metabolites produced in incubations with mouse liver microsomes when chloroform concentrations are not too high (Vittozzi et al., 1991). The relative importance of the minor pathways of phosgene metabolism depends upon the availability of glutathione, other thiols, and other nucleophilic compounds, such as histidine and cysteine (see Figure 2).

Fig. 2: Metabolism of chloroform

(GSH = glutathione; GSSG = bis(gamma-glutamyl-L-cysteinylglycine) disulfide; Nu = tissue nucleophiles; R = alkyl group)

Oxidative metabolism, with CYP2E1 (an ethanol-inducible mono-oxygenase isoenzyme system present in the liver of mammals, including humans) playing a key role, is probably the only significant in vivo pathway at low exposures, and available data indicate that oxidative metabolism has a major role in toxicity. The dominant role of CYP2E1 in metabolizing chloroform to toxic metabolites has been demonstrated in studies involving treatment of animals with enzyme inducers or inhibitors, as well as studies in mice lacking CYP2E1 (Brady et al., 1989; Guengerich et al., 1991; Nakajima et al., 1995a,b; Constan et al., 1999; see also section 8.8). Immunoinhibition studies with anti-CYP2E1 monoclonal protein have shown that CYP2E1 is responsible for 81% of the metabolism assayed at a low chloroform (0.5 mmol/litre) concentration in liver microsomes from acetone-induced rats (Brady et al., 1989). Toxicity to rat and mouse hepatocytes incubated in vitro with up to 5 mmol chloroform/litre was prevented by the addition of a CYP2E1 inhibitor or by reduced oxygen tension, underscoring the importance of oxidative metabolism in toxicity (Ammann et al., 1998). Regional distribution of liver lesions in rats and mice correlates well with the hepatic distribution of CYP2E1 and glutathione (Smith et al., 1979; Ingelman-Sundberg et al., 1988; Tsutsumi et al., 1989; Johansson et al., 1990; Dicker et al., 1991; Nakajima et al., 1995a,b).

CYP2B1 may also have a role in chloroform metabolism, although this is likely to be only minor at low tissue chloroform concentrations (studies reviewed in Environment Canada & Health Canada, 2001). However, at high tissue concentrations (e.g., resulting from an oral dose of 0.5 ml/kg body weight), chloroform hepatotoxicity was dramatically potentiated in Wistar rats treated with phenobarbital (a CYP2B1 inducer) but not in rats treated with n-hexane (a CYP2E1 inducer), compared with uninduced controls (Nakajima et al., 1995b).

A study in which rats were exposed to [¹⁴C]chloroform showed that metabolism was most active in the liver, followed by the nose and kidney. Metabolic activity was correlated with accumulation of metabolites (Löfberg & Tjälve, 1986).

The first extensive PBPK model for chloroform described liver and kidney individually as metabolic sites for chloroform. The maximum velocity of metabolism in the kidney was scaled to that in the liver, and terms were introduced to account for loss and resynthesis of metabolizing enzyme (Corley et al., 1990). This model was modified to include a description of liver cytotoxicity (Reitz et al., 1990). Later, Gearhart et al. (1993) modified the tissue:blood partition coefficients according to temperature and fitted gas uptake without the need to describe enzyme loss and resynthesis. Others subsequently incorporated absorption from the stomach as well as the gastrointestinal tract and also accounted for gastric emptying time (Dix et al., 1994; Dix & Borghoff, 1995). In 1996, kidney and liver model compartments were subdivided into regions of high and low metabolic activity (Lilly, 1996). The combination of this approach with the two-compartment absorption model of Dix & Borghoff (1995) resulted in a recent PBPK model in the "hybrid" species⁵ (ILSI, 1997).

Health Canada developed a model for the dog, using physiological and anatomical parameters from Brown et al. (1997), while metabolic parameters were based on the average of rat and human parameters. The fractional subvolumes for the liver were assumed to be the same as those reported for the rat by ILSI (1997) (Environment Canada & Health Canada, 2001).

Health Canada also developed a human model. Physiological parameters were derived from Brown et al. (1997), with the exception of the ventilation rate and cardiac output, which were related to an assumed breathing rate of 23 m³ air/day. ILSI (1997) was used as the source of the partition coefficients and rate constants. Liver tissue subvolumes were assumed to be the same as in the rat, while kidney was subdivided into a 70:30 cortex:non-cortex ratio. Human metabolic parameters had been determined in vitro in eight human liver samples, as reported by Corley et al. (1990). Kidney rate constants were based on the relationship of activity observed in the microsomal fraction of kidneys to the activity observed in the microsomal fraction of the liver, based on the in vitro results reported by Corley et al. (1990), but supported by data on metabolism of two known substrates of CYP2E1 by microsomal fractions of the kidney and liver from 18 humans, reported by Amet et al. (1997). As it is based on metabolized dose, the model accounts for differences in metabolism between humans and (in this case, hybrid) laboratory animals (Environment Canada & Health Canada, 2001).

Results from the human model were compared with data on total metabolized parent and exhaled chloroform reported by Fry et al. (1972), where chloroform was administered, in olive oil or gelatin capsules, to male and female volunteers. Exhaled chloroform was measured for up to 8 h following dosing, and the total percentage of the dose exhaled unchanged was calculated by extrapolation to infinite time. Human model simulations conducted using a single-compartment description of oral uptake were closer to the observations of Fry et al. (1972) than those estimated using a multi-compartment description. Therefore, while a multi-compartment description was necessary in the rat model, a single-compartment description of oral uptake was used in estimating human-equivalent concentrations (Environment Canada & Health Canada, 2001). The model was also modified to permit assessment of exposure to chloroform from all likely sources, including air, water, and food. The exposure scenario (see section 11.1.3) was modelled within a 24-h day and included inhalation, ingestion, and dermal exposure from one 10-min shower, a brief washing-up period before retiring at night, discrete periods of food and water consumption, and inhalation of chloroform at various concentrations (ICF Kaiser, 1999; Environment Canada & Health Canada, 2001).

The physiological and metabolic parameter values for rats, dogs, and humans used to exercise the PBPK model are reproduced in Table 5. In the liver, the V_max for the metabolism of chloroform is twice as high in humans as in rats, while there is little difference in V_max in the kidney or in the affinity (K_m) in either liver or kidney (see Table 5). Further details are available in the source document (Environment Canada & Health Canada, 2001).

Table 5: PBPK model physiological and metabolic parameter values in rats, dogs, and humans.

Chloroform has a moderate acute oral toxicity in rats, with LD₅₀s ranging from 0.45 to 2.0 g/kg body weight (Kimura et al., 1971; Chu et al., 1980). In mice, a wide range (36-1366 mg/kg body weight) of acute oral LD₅₀ values has been reported (IPCS, 1994a). Acute oral administration produced narcosis and anaesthesia in rodents (IPCS, 1994a). An increase in renal cell proliferation was seen in male Osborne-Mendel and F344 rats given gavage doses of 10 and 90 mg/kg body weight, respectively (Templin et al., 1996a). In male F344 rats, a no-observed-adverse-effect level (NOAEL) and a lowest-observed-adverse-effect level (LOAEL) for serum enzyme changes indicative of liver damage following acute gavage exposure have been established as 30 and 60 mg/kg body weight, respectively (Keegan et al., 1998). Administration of 0, 67, 135, or 338 mg/kg body weight by gavage in olive oil to male Wistar rats increased, in a dose-dependent manner, the number of necrotic hepatocytes in the centrilobular region and elevated plasma alanine aminotransferase (ALAT) levels significantly (Nakajima et al., 1995b). Liver and kidney changes were seen in rats administered chloroform at 250 mg/kg body weight by gavage (Torkelson et al., 1976). Cell proliferation occurred in the liver and kidneys of male B6C3F1 mice given 150 mg/kg body weight by gavage; severe necrosis was also seen in the kidneys (Gemma et al., 1996). Hepatic necrosis was observed in male mice 48 h following a single gavage administration of 240 mg/kg body weight (Reitz et al., 1982). Minimal centrilobular enlargement was observed in male mice 4 days following intragastric administration of 66 mg/kg body weight (Moore et al., 1982). Lesions and epithelial cell proliferation were seen in the nasal passage of F344 and Osborne-Mendel rats following gavage administration of 90 mg/kg body weight in corn oil (Templin et al., 1996a).

An inhalation LC₅₀ value (for 6-h exposure) of 9.2 g/m³ has been reported in rats (Bonnet et al., 1980). No deaths occurred when F344 rats (10 per sex per concentration) were exposed for 6 h at up to 5 g/m³, but 17/20 died at 10 g/m³ (Kasai et al., 2002). Depression of the central nervous system is a dominant symptom of acute inhalation. Rats exposed at 2.1 g/m³ for 4 h showed significant subnarcotic effects (Frantík et al., 1998).

In female OF1 mice, an inhalation LC₅₀ value (for 6-h exposure) of 6.2 g/m³ was reported (Gradiski et al., 1978). Groups of 10 female BDF1 mice survived a 6-h exposure at up to 2.5 g/m³, but died (showing centrilobular liver necrosis) at 40 g/m³. Male mice are much more susceptible to acute chloroform inhalation toxicity, 1 of 10 and 8 of 10 dying after a single 6-h exposure at 59 and 120 mg/m³, respectively. The cause of death in the males was necrosis of the proximal tubules of the kidneys (Kasai et al., 2002).

Kidney tubule degeneration was seen in rabbits following a 24-h covered application of 1 g/kg body weight; no gross changes were seen in the liver (Torkelson et al., 1976).

8.2.1 Ingestion

Lesions and cell proliferation in the olfactory epithelium and changes in the nasal passages were seen in female F344 rats given 34 mg/kg body weight per day for 4-5 days in corn oil by gavage; after 3 weeks of administration, these effects were observed at 100 but not at 34 mg/kg body weight per day (Larson et al., 1995a; Dorman et al., 1997).

Tissue changes in the kidneys (mineralization, hyperplasia, and cytomegaly) and liver (inflammation) were seen in mice given 37 mg/kg body weight per day by gavage for 14 days (Condie et al., 1983).

8.2.2 Inhalation

F344 rats and BDF1 mice (10 per sex per species per concentration) were exposed at 0, 2.5, 5, 10, 20, or 40 g/m³, 6 h/day, 5 days/week, for 2 weeks. Rats survived at up to 5 g/m³, but all died within 2 days at 10 g/m³ and above. The female mice survived exposure at 2.5 g/m³, but deaths occurred (from day 4 onwards) at 5 g/m³. Only two male mice survived, one at 2.5 g/m³ and one at 5 g/m³. Dead rats showed lung congestion and inflammation, believed to arise as a result of cardiovascular toxicity. Deaths in mice were ascribed to necrosis of the kidney proximal tubules (males) and centrilobular necrosis of the liver (females). Surviving rats showed vacuolic changes in the proximal kidney tubules and the central area of the liver, as well as desquamation, atrophy, and disarrangement of the olfactory epithelium and oedema of the lamina propria of the nasal cavity. Surviving male mice had necrosis in the kidney proximal tubules, slight swelling and vacuolic change in the liver, and atrophy and respiratory metaplasia in the olfactory epithelium. The surviving female mice showed necrosis and vacuolic changes in the liver and necrosis and disarrangement of the olfactory and respiratory epithelia, but no kidney changes (Kasai et al., 2002).

Cell proliferation was observed in the nose (ethmoid region) of male F344 rats following inhalation exposure at 9.8 mg/m³ for 6 h/day for 4 days. At 49 mg/m³, only minimal to mild lesions were seen (Templin et al., 1996b). Following exposure to 50 mg/m³ for 6 h/day for 7 consecutive days, male F344 rats had lesions in nasal turbinates, including increased cell proliferation in central, proximal, and distal regions of the first endoturbinate, and histological changes in the central turbinate bone (Larson et al., 1994a; Mery et al., 1994).

Cell proliferation was seen in the nasal turbinates of female B6C3F1 mice exposed at 10 mg/m³, but not at 1.5 mg/m³, for 6 h/day, 7 days/week, for 3 weeks (Larson et al., 1996). Increased cell proliferation was detected in the first endoturbinate of the nasal passage in female B6C3F1 mice exposed to 49 mg chloroform/m³, 6 h/day, for 7 consecutive days (Mery et al., 1994). No microscopic damage was seen in the nasal passages of female B6C3F1 mice exposed to up to 1500 mg/m³ for 6 h/day for 7 consecutive days. Cell proliferation was not measured (Larson et al., 1994a).

Studies designed specifically to investigate cytotoxicity and regenerative cell proliferation in target organs are mentioned briefly (without experimental details) in section 8.8.

Studies designed specifically to investigate cytotoxicity and regenerative cell proliferation in target organs are mentioned briefly (without experimental details) in section 8.8.

8.3.1 Ingestion

When B6C3F1 mice were given 60 mg/kg body weight per day and above (the other doses were 130 and 270 mg/kg body weight per day) for 90 days by gavage in corn oil, both sexes showed increased liver weights and vacuolation and lipid accumulation in the liver. When chloroform was given in an Emulphor vehicle, the only effect at the lowest dose level (60 mg/kg body weight per day) was increased liver weight in females. No chloroform-related histopathological changes were observed in the kidneys (Bull et al., 1986).

In a study in which groups of 7-12 male and female CD1 mice were given 0, 50, 125, or 250 mg/kg body weight per day by stomach tube for 90 days, effects noted at all doses were increased liver weight and increased hepatic microsomal activity (in females) and (in both sexes) microscopic tissue changes in the liver (hepatocyte degeneration and focal lymphocyte collection) and kidneys (intertubular collection of inflammatory cells) (Munson et al., 1982).

When female B6C3F1 mice were exposed to chloroform in drinking-water for 90 days, fatty changes in the liver were observed at 263 mg/kg body weight per day (US EPA, 1980).

In male Osborne-Mendel rats, liver cholesterol was significantly increased when chloroform was given at 81 mg/kg body weight per day for 90 days in drinking-water (US EPA, 1980).

Effects on the liver were reported when chloroform was administered at 15 or 30 mg/kg body weight per day in a toothpaste base in gelatin capsules to male and female beagle dogs for 6 days/week for 7.5 years. Sacrifice followed a subsequent period of 19-23 weeks without treatment. The protocol included vehicle controls, untreated controls, and dogs receiving alternative (non-chloroform) toothpaste. Each group contained 8 animals of each sex, with the exception of the vehicle control group, which included 16 of each sex. At the high dose, there were significant increases in serum ALAT levels at 6 weeks of treatment. At the low dose, significant increases in ALAT levels were observed at 34 weeks and later. "Fatty cysts" of the liver were observed in both dose groups at the end of the study. In males, the incidences of fatty cysts of moderate or marked severity were 1/15, 6/7, and 6/7 in the vehicle control, low-dose, and high-dose groups, respectively. Equivalent incidence figures in the females were 0/12, 3/8, and 7/8 for the vehicle control, low-dose, and high-dose groups, respectively (see Table 6). There were no treatment-related increases in tumours (Heywood et al., 1979). The fatty cysts might reflect chronic low-grade disruption of hepatocyte function. Hepatocytes play a vital role in synthesis and transport of lipoprotein, triglyceride, and fatty acid metabolism. These fatty cysts (granulomas) are commonly seen in old dogs and probably form after the rupture or fusion of fat-laden cells. The subsequent macrophage response results in foamy aggregates in sinusoids, portal triad stroma, and hepatic venules, possibly accompanied by multinucleate giant cells and ceroid pigment accumulation resulting from lipid breakdown (D. Malarkey, personal communication to Health Canada, 2003). [LOAEL = 15 mg/kg body weight per day]

Table 6: Fatty cyst incidence in dog study.^a

^a From Heywood et al. (1979).

8.3.2 Inhalation

Exposure of rats (10-12 per sex per exposure level; strain unspecified) at 120, 240, or 420 mg/m³, 7 h/day, 5 days/week, for 6 months resulted in increased relative kidney weight, cloudy swelling in the renal tubular epithelium, and focal liver necrosis in the males at all dose levels (Torkelson et al., 1976).

Nasal effects were seen in a study in which female B6C3F1 mice were exposed at 1.5, 9.8, 49, 147, or 441 mg/m³, 6 h/day, for up to 13 weeks. After 3 weeks, there was slightly increased proliferation in the nasal turbinate lamina propria at 9.8 mg/m³ and above and mild to minimal nasal lesions at 49 mg/m³ and above. After 13 weeks, nasal effects were minimal at 49 and 147 mg/m³, but cell proliferation persisted at 441 mg/m³ (Larson et al., 1996). [LOAEL = 49 mg/m³]

When BDF1 mice (10 per sex per exposure level) were exposed by inhalation at 0, 59, 123, 245, 490, or 980 mg/m³ for 6 h/day, 5 days/week, for 13 weeks, the females all survived, but reduced growth and deaths occurred in males in all chloroform groups. Increased liver and kidney weights occurred in the mice at 980 mg/m³. Male mice showed effects in the kidney (necrosis of the proximal tubules) and nasal cavity (bone thickening and degeneration of the olfactory epithelium) at all exposure levels (59 mg/m³ and above). Female mice showed nasal cavity toxicity (thickening of bone, eosinophilic changes in the olfactory and respiratory epithelia) at all exposure levels. The liver was normal in both sexes at up to 245 mg/m³, but cell atypia was seen in the females at 490 mg/m³, and swelling or necrosis occurred at 980 mg/m³ (Kasai et al., 2002). [LOAEL = 59 mg/m³]

Daily exposure of male F344 rats for 6 h/day for 13 weeks resulted in mild histological changes in nasal passages at 9.8 mg/m³ and increased cell proliferation with enhanced bone growth at 49 mg/m³. At 147 mg/m³ and above, the cortical proximal tubules of the kidneys showed increased epithelial cell proliferation. Hepatic lesions (including cell proliferation) were seen at 1470 mg/m³ (Templin et al., 1996b). [LOAEL = 9.8 mg/m³]

When F344 rats (10 per sex per exposure level) were exposed by inhalation at 0, 123, 245, 490, 980, or 1960 mg/m³ for 6 h/day, 5 days/week, for 13 weeks, growth was reduced and liver and kidney weights were increased at 245 mg/m³ and above. The nasal cavity tissues were the most sensitive target, showing mineralization and atrophy of the respiratory epithelium at all exposure concentrations and necrosis at 980 mg/m³ and above. The liver and kidneys were microscopically normal at up to 245 mg/m³, but there were vacuolic changes in the kidneys and liver cell collapse/hepatocyte loss at 490 mg/m³ and above (Kasai et al., 2002). [LOAEL = 123 mg/m³]

8.4.1 Overview of carcinogenicity

Chloroform induced liver tumours in mice (both sexes) when given by gavage in a corn oil vehicle and possibly in males when given (orally) in a toothpaste vehicle, but was not carcinogenic to the mouse liver when given in the drinking-water or by inhalation. Chloroform induced kidney tumours in male mice exposed by inhalation or (in one of four strains) by ingestion in a toothpaste vehicle, but not when given in corn oil. Female mice were not susceptible to chloroform-induced kidney cancer.

Chloroform induced kidney tumours in one strain of male rats treated via a corn oil vehicle and in drinking-water, but not in another strain treated via the drinking-water. Chloroform was not carcinogenic in inhalation studies or where given (orally) in a toothpaste base. Female rats were not susceptible to chloroform-induced kidney cancer. Chloroform did not induce liver tumours when given in corn oil or by inhalation, although two drinking-water studies found limited evidence of liver carcinogenicity, one in females only and a second in which only males were treated.

There were no clear descriptions of non-neoplastic lesions in the tumour target organs in the reports of these studies; consequently, a phase of research work has since been undertaken to investigate early changes in the target organs at the doses that were active in the chronic studies. In addition, re-evaluation of the kidney tissue slides has been undertaken for the Jorgenson et al. (1985) study. There has also been a very limited re-evaluation of the kidney tissues from the male rats of the NCI (1976) study.

8.4.2 Liver

Chloroform induced liver tumours in male and female B6C3F1 mice following gavage at 138 mg/kg body weight per day or more in a corn oil vehicle for 78 weeks (NCI, 1976), but not when given (to females only) at up to 1800 mg/litre in the drinking-water for 104 weeks (up to 263 mg/kg body weight per day) (Jorgenson et al., 1985). It did not significantly increase the incidence of liver tumours when given by gavage (in a toothpaste) at up to 60 mg/kg body weight per day, 6 days/week, for 80 weeks to ICI mice (both sexes) or male C57BL, CBA, or CF1 mice (Roe et al., 1979). Chloroform induced a statistically borderline significant increase in hepatic tumours (adenomas and carcinomas combined) in female but not in male BDF1 mice exposed by inhalation at up to 441 mg/m³, 6 h/day, 5 days/week, for 104 weeks (Yamamoto et al., 2002).

Chloroform did not induce liver tumours when given by gavage to male or female Osborne-Mendel rats at up to 200 mg/kg body weight per day for 111 weeks (NCI, 1976), in the drinking-water at up to 160 mg/kg body weight per day to male Osborne-Mendel rats (Jorgenson et al., 1985), orally in a toothpaste at up to 165 mg/kg body weight per day for 80 weeks to male or female Sprague-Dawley rats (Palmer et al., 1979), or by inhalation at up to 441 mg/m³ in male or female F344 rats exposed for 6 h/day, 5 days/week, for 104 weeks (Yamamoto et al., 2002). In a study in which Wistar rats were treated (at >100 mg/kg body weight per day for up to 185 weeks) via the drinking-water, males did not develop liver tumours, but a statistically significant increase in foci of cellular atypia (possibly representing pre-neoplastic lesions) was seen in the females. However, the relevance of this finding is difficult to determine because the control group was small (22) and the exposed females survived longer (185 weeks) than the controls (145 weeks), so there was no basis for comparing the incidence of late-developing tumours (Tumasonis et al., 1987). According to a recent abstract, administration of chloroform at 800 mg/litre in the drinking-water for 100 weeks did not induce liver cancer in a group of 78 male F344 rats. However, at the top dose (1600 mg/litre water; probably about 160 mg/kg body weight per day), the incidence of rats with hepatocellular neoplasia (adenoma or carcinoma) was increased (17% versus 5.1%; P < 0.05). The significance of this result cannot be clearly determined from the limited presentation of the data (DeAngelo et al., 2003).

No evidence of carcinogenic potential was seen in a study in which male and female wild-type and rasH2-Tg transgenic CB6F1 mice were given up to 140 (males) or 240 (females) mg/kg body weight, 5 days/week for 26 weeks, by gavage in corn oil. A range of organs was examined, including the liver. Swelling and vacuolation of hepatocytes were seen in the transgenic and non-transgenic mice, and the incidence of hepatocellular foci was increased in the female mice (transgenic and wild type) at the top dose (Sehata et al., 2002). In another study in which hemizygous transgenic (Tg.AC) mice were given the same doses of chloroform by gavage in corn oil, 5 days/week for 13 weeks, V-Ha-ras transgene expression was not induced in the liver, although there was evidence of liver tissue damage and cell proliferation (Delker et al., 1999). (These transgenic mouse model studies were primarily conducted for testing their suitability/validity as a bioassay for showing tumorigenicity of a positive "non-genotoxic-cytotoxic rodent liver and kidney carcinogen" using chloroform as a prototype model carcinogen of this category of chemical carcinogens.)

8.4.3 Kidney

Chloroform induced kidney tubular cell carcinoma in male BDF1 mice treated by inhalation at 147 or 441 mg/m³, 6 h/day, 5 days/week, for 104 weeks. No clear response was seen at 25 mg/m³, the lowest exposure concentration (Yamamoto et al., 2002). Chloroform induced a mixture of malignant hypernephromas and benign cortical adenomas in the kidneys of male ICI mice given the compound at 60 mg/kg body weight per day for 80 weeks in a toothpaste vehicle (Roe et al., 1979). It did not induce kidney tumours in B6C3F1 mice (either sex) treated at up to 477 mg/kg body weight per day for 78 weeks by corn oil gavage (NCI, 1976), in female B6C3F1 mice ingesting up to 263 mg/kg body weight per day for 104 weeks via the drinking-water (Jorgenson et al., 1985), in female BDF1 mice exposed by inhalation at up to 441 mg/m³ for 104 weeks (Yamamoto et al., 2002), or in female ICI or male C57BL, CBA, or CF/1 mice given chloroform at 60 mg/kg body weight per day for 80 weeks in toothpaste (Roe et al., 1979).

Chloroform induced kidney epithelial tumours in male Osborne-Mendel rats treated at 180 mg/kg body weight per day for 111 weeks by gavage in corn oil (NCI, 1976) and kidney tubular cell adenomas and adenocarcinomas in male rats of the same strain given drinking-water containing chloroform at 1800 mg/litre (providing 160 mg/kg body weight per day) for 104 weeks (Jorgenson et al., 1985; see Table 7). It did not cause kidney cancer in male F344 rats when given at up to 1600 mg/litre in the drinking-water (probably about 160 mg/kg body weight per day) for 100 weeks (DeAngelo et al., 2003), in female Osborne-Mendel rats when given at up to 200 mg/kg body weight per day for 111 weeks in corn oil (NCI, 1976), in male or female F344 rats exposed by inhalation at up to 441 mg/m³, 6 h/day, 5 days/week, for 104 weeks (Yamamoto et al., 2002), in male or female Wistar rats treated at up to 2900 mg/litre drinking-water (average doses about 180-240 mg/kg body weight per day) for up to 185 weeks (Tumasonis et al., 1987), or in male or female Sprague-Dawley rats given chloroform at up to 165 mg/kg body weight per day for 80 weeks in toothpaste (Palmer et al., 1979).

Table 7: Histopathology in kidneys of male Osborne-Mendel rats in drinking-water bioassay of Jorgenson et al. (1985).^a

^a From Hard et al. (2000).

^b Lesions indicative of tubule injury included nuclear crowding, cytoplasmic vacuolation, and faint basophilia in the mid to deep cortex.

In a study identifed only as an abstract, transgenic p53^+/- (sensitive to mutagenic carcinogens) and p53^+/+ wild-type mice were given chloroform at up to 140 or 240 mg/kg body weight per day (males and females, respectively) by gavage in corn oil for up to 26 weeks. The males showed renal tubular regeneration and proliferation, but there were no increases in kidney tumour incidence (Gollapudi et al., 1999). No evidence of carcinogenic potential was seen in another study in which male and female wild-type and rasH2-Tg transgenic CB6F1 mice were given chloroform at up to 140 (males) or 240 (females) mg/kg body weight, 5 days/week for 26 weeks, by gavage in corn oil. A range of organs was examined, including the kidneys (Sehata et al., 2002). In another study in which hemizygous transgenic (Tg.AC) mice were given the same doses of chloroform by gavage in corn oil, 5 days/week for 13 weeks, V-Ha-ras transgene expression was not induced in the kidneys, although there was evidence of kidney tissue damage and cell proliferation in the males (Delker et al., 1999).

The bioassay in which evidence of carcinogenicity was seen at the lowest concentration or dose, and which involved exposure similar to that of humans, was that of Jorgenson et al. (1985). Male Osborne-Mendel rats were given chloroform in the drinking-water at 0, 200, 400, 900, or 1800 mg/litre for 2 years. Group sizes were 330, 330, 150, 50, and 50 respectively, and the TWA daily doses were